Global change (climate change together with other worldwide anthropogenic processes such as increasing trade, air pollution and urbanization) will affect plant health at the genetic, individual, population and landscape level. Direct effects include ecosystem stress due to natural resources shortage or imbalance. Indirect effects include (i) an increased frequency of natural detrimental phenomena, (ii) an increased pressure due to already present pests and diseases, (iii) the introduction of new invasive species either as a result of an improved suitability of the climatic conditions or as a result of increased trade, and (iv) the human response to global change. In this review, we provide an overview of recent studies on terrestrial plant health in the presence of global change factors. We summarize the links between climate change and some key issues in plant health, including tree mortality, changes in wildfire regimes, biological invasions and the role of genetic diversity for ecosystem resilience. Prediction and management of global change effects are complicated by interactions between globalization, climate and invasive plants and/or pathogens. We summarize practical guidelines for landscape management and draw general conclusions from an expanding body of literature.

A growing body of literature is addressing the potential impacts of global change on plant pathosystems. Global change encompasses changes in climate, connectivity among continents, concentration of air pollutants and other pervasive processes such as urbanization. All these processes and their interactions are predicted to affect plant health at the genetic, individual, population and ecosystem level. Plant health is the outcome of the dynamic interactions of a diversity of plant hosts with a similar wide spectrum of plant pathogens under a variety of environmental conditions. Plant health is not only related to the presence and severity of plant diseases and pests, but can also be impaired by abiotic disturbances and a range of stress factors. There is thus a need to broaden the traditional definition of plant health, from the study of specific pathosystems in controlled conditions to broad and interconnected issues at the landscape and species distributional range level (e.g. Gäumann, 1948; Tarr, 1972; Robinson, 1976; Zadoks & Schein, 1979; Geils, 1992; Ferguson, 1994; Tainter & Baker, 1996; Holdenrieder et al., 2004; Madden, Hughes & Van den Bosch, 2007; Money, 2007; Ostry & Laflamme, 2009; Fig. 1).

Figure 1. Conceptual overview of this review. Global change will affect plant health through direct effects on hosts (e.g. host debilitation owing to climate change, pollution, etc.) and indirect effects due to changed biotic and abiotic conditions. Landscape management can have an influence on both plant health and global change by acting on hosts and on the various stressors (plant pathogens, pests, wildfires, windstorms and droughts). Global change factors (climate change, increased trade, land-use change, pollution and urbanization) are interrelated too. To be successful in the long term, landscape management in the face of global change will need an adaptive strategy. Numbers in parenthesis refer to sections in this review.

The implications of global change for landscape management and the rural economy have only rarely been addressed. Compared to the number of reviews on the potential effects of global change on plant pathogens and pests (Table 1), there are still relatively few studies demonstrating such effects. More research has modeled long-term scenarios of plant epidemic development in the presence of climate change. These studies and reviews still need to be put in perspective, given that global change will not only imply a change in temperature and precipitation patterns, but also other processes such as host range expansion and pathogen introductions. One of the main points of this review is that studies of host-pathogen systems are likely to fail in predicting future plant health if they overlook the contribution to pathosystems of global change factors such as increased trade and other stressors. Here, we aim to discuss selectively the existing body of literature on plant health and global change and the evidence base for decision-making on invasive plant pathogens and exotic plants. Wherever possible, we try to point out the practical implications of these studies for a sustainable vegetation management in heterogeneous landscapes.

Table 1. Selected reviews relating to global change and plant health

Main focus

Selected conclusion

Reference

O3, SO2 and acid rain

“Depending on the particular pollutant/host/pathogen interaction, there may be either an increase, decrease, or no change in disease development”

“Climate change-induced modifications of frequency and intensity of forest wildfires, outbreaks of insects and pathogens, and extreme winds, may be more important than the direct impact of higher temperatures and elevated CO2.”

“It is extremely important for forest managers to begin to include climate considerations in their strategic and operational plans yet, to date, there is little evidence that many are taking a proactive approach to the issue”

Research gaps include “multiple environmental stresses in critical load determinations; interactions between air pollution, climate, and forest pests; effects of forest fire on air quality; and forest carbon sequestration under changing climate and co-exposure to elevated levels of air pollutants”

Disturbance processes such as fire, insect outbreaks, plant diseases and water flow affect landscapes, including forests, savannas and grasslands, at spatial scales over hundreds of kilometres to time scales of years to decades (Holling, 1992). A challenge is to predict the impact of climate change, specifically of global warming, on the occurrence and severity of these processes and the likely success of adaptation and mitigation options. Disease can have a profound impact on landscape management. For example, the crisis in the agricultural economy following the Black Death pandemic in 14th Century Europe, together with political instability and climate deterioration, is believed to have led to general abandonment of marginal arable land, a decrease in grazing pressure, and an increase in forest land cover (van Hoof et al., 2006; Yeloff & van Geel, 2007; Pongratz et al., 2008). The epidemics accompanying the European conquest of America had a similar outcome on land cover through a diminution of fire use by the dwindling indigenous population (Nevle & Bird, 2008). Due to climate change, it is likely that some regions may experience future climate conditions never experienced before, and this will present an additional challenge for a sustainable land use management (Williams, Jackson & Kutzbach, 2007). New climate conditions might lead not only to unexpected species associations but also to the disruption of traditional land use patterns.

In the rest of this review, we discuss broad issues related to plant health and global change, including: (i) recent studies linking climate change with tree mortality rates. One of the ways tree mortality rates will be affected by climate change is through (ii) changes in wildfire regimes. This is also true for tropical ecosystems, although there is a relative dearth of studies assessing (iii) the importance of global change for the tropics. A similar relative lack of research applies for (iv) regional studies of tree genetic diversity in the context of climate change (genetic diversity as an insurance policy); there is a need for this research direction to move from its current historical perspective (what happened during the last glaciations) to a more applied approach (how resilient are current tree populations to future climate change). Climate change will not only affect forested landscapes, but will also be relevant for (v) the health and management of grassland ecosystems. For forests, grasslands and other habitats, there is the additional complication of (vi) the interactions between globalization, climate change and biological invasions. We draw general conclusions from an expanding body of literature and summarize practical guidelines for landscape management.

Forests are believed to be particularly susceptible to abrupt climate change because of their longevity (e.g. Cannell, Grace & Booth, 1989). Forests are also one of the highest sources of uncertainty in the response of the biosphere to increased levels of CO2 in the atmosphere (Purves & Pacala, 2008). Thanks to their long age, trees do generally have some capacity to withstand variations in environmental conditions (Hamrick, 2004; Petit & Hampe, 2006; see also Phillips, Buckley & Tissue, 2008), but if these conditions move rapidly beyond the range typically encountered by a certain tree species, population, or individual (e.g. Overpeck & Cole, 2006), the long periods necessary to reconstitute a mature forest stand imply a particular vulnerability of trees to climate change, as shown by studies of the impacts on terrestrial ecosystems of El Niño events (Holmgren et al., 2001). In this section, we discuss recent studies linking tree mortality rates with climatic changes, with a particular focus on temperate forests, although similar issues apply also to other biomes.

There is evidence that changes in precipitation might already be increasing rates of tree mortality in some regions. For example, the mortality rate of trees in old-growth forest plots in the Sierra Nevada, California, was found to increase over the period 1983–2004. This increase was not paralleled by a similar increase in recruitment rates, but was associated with an increase in drought (van Mantgem & Stephenson, 2007). Analogous findings have now been reported for a wider region on the West Coast of the USA (van Mantgem et al., 2009). Similarly, an increasing water stress effect on radial growth of Abies alba during the second half of the 20th Century was reported from forests at the South-Western margins of the distributional range of this drought-sensitive tree species (Macias et al., 2006). An analogous case study is found in an inner Swiss valley, the Valais, where Scots pine is experiencing widespread dieback and is believed to be already threatened by climate change. Tree defoliation levels were found to be negatively correlated with the precipitation in the previous year (Rebetez & Dobbertin, 2004), and this is a worrisome finding for the future, given that the climate in that region is shifting towards longer summer drought periods (Weber, Bugmann & Rigling, 2007).

In regions where temperature, but not precipitation, is predicted to increase, these and other water-limited forests are likely to experience considerable abiotic stress (e.g. Seidling, 2007; McDowell et al., 2008). Such disturbance is often followed by an increased activity of secondary pests and pathogens (Schwartz, 1992; Brasier & Scott, 1994; Desprez-Loustau et al., 2006b), with subsequent potentially extensive dieback at the landscape scale (Holdenrieder et al., 2004). This can in turn lead to an enhanced habitat for saproxylic organisms which require an amount of coarse and fine woody debris larger than what is normally found in today's managed forested landscapes (Lonsdale, Pautasso & Holdenrieder, 2008). In some cases, forests, tree species and individual trees may be resistant to occasional drought years, but may be susceptible to multi-year droughts, as shown (i) by Scots pine decline in 20th Century Valais (Bigler et al., 2006), (ii) by a historical analysis of the effects of a persistent drought event (1922–1932) in Inner Mongolia on the mortality of Picea koraiensis and Pinus tabulaeformis (Liang et al., 2003), and (iii) by episodic mortality of long-lived desert shrubs in the Colorado Desert of California (Miriti et al., 2007). In other cases, tree mortality due to strong droughts may be higher for tree individuals already weakened by droughts in previous decades. Evidence for such a long-term signal comes from a study of the effects of a strong La Niña event on the mortality of Nothofagus dombeyi in Patagonia (Suarez, Ghermandi & Kitzberger, 2004), and from an analysis of the effects of the 1994 drought on Quercus ilex forests in Catalonia (Lloret, Siscart & Dalmases, 2004).

Trees are thus likely to be particularly sensitive to changes in the frequency and severity of extreme events. Evidence is reported from subalpine forests in the Rocky Mountains, where early-season and late-season drought resulted in an increase in the mortality of Picea engelmannii and of Abies lasiocarpa, but not of Pinus contorta (Bigler et al., 2007). Interspecific differences in the effects of climate change on tree mortality and recruitment rates are expected to lead to changed patterns of competition, thus potentially causing shifts in the geographic mosaic of species composition and co-evolution. Similarly, climate change might differentially affect different cohorts of a certain species. There is evidence that this may be the case in pinyon-juniper woodland of Northern Arizona, where drought years increased the mortality of large individuals of Pinus edulis relative to small ones and also relative to Juniperus monosperma (Müller et al., 2005). Past climatic changes have played an important role in the expansion and contraction of the pinyon-juniper vegetation of the Western US, with pinyon expanding in wetter periods and juniper in drier ones, and both species retreating during colder spells (Romme et al., 2009).

Different effects of climate change on tree growth and survival can occur for populations of the same species from different parts of their distributional range. This is shown by climate gradient experiments for Pinus sylvestris populations in Europe and North America (Reich & Oleksyn, 2008), where differences among populations were related to climate transfer distance, with enhanced versus diminished growth and survival in the northern versus southern parts of the species range. An analogous shift in tree species distributions can occur at different elevations, as climate change is likely to lead to different patterns of plant recruitment and loss at different levels of an altitudinal gradient, as shown by the declining populations of Chamaecyparis nootkatensis at low altitudes in south-eastern Alaska (Beier et al., 2008). Another example of the differential response of a tree species to climate change at different altitudinal levels is provided by a study of the influences of climate on the population dynamics of Pinus jeffreyi in the Carson Range, Nevada, with a retracting population at low altitudes and an expanding population at middle and, more slowly, at high altitudes (Gworek, Wall & Brussard, 2007).

A challenge in predicting intra- and inter-specific changes due to climate change will be disentangling the effects of overall trends from those of extreme years. The regular, long-term cycle of larch bud moth (Zeiraphera diniana) outbreaks in the Alps (Baltensweiler, Weber & Cherubini, 2008; Buntgen et al., 2009) was interrupted in 1990 by unusually cold spring weather conditions (Baltensweiler, 1993). In the Eastern and Northern Iberian Peninsula, an increased variability in precipitation due to a higher occurrence of extreme years was linked to an increased variability in tree growth for Pinus nigra, P. sylvestris and P. uncinata (Andreu et al., 2007). But, at the same time, the shared variability among the tree chronologies, the frequency of narrow rings and the year-to-year growth variability increased for the period 1885–1992 (Andreu et al., 2007). A more synchronous growth among tree species might be an indication for climate having become more limiting to growth. In Northern Patagonia, tree mortality of Austrocedrus chilensis, a xeric conifer, is intensified by extreme events of the El Niño Southern Oscillation (Villalba & Veblen, 1998). But, at the same time, variations in the intensity of droughts in different regions are associated with a latitudinal variability in tree mortality (Villalba & Veblen, 1998). Additional complications are the contributions of a newly discovered root pathogen, Phytophthora austrocedrae (Greslebin, Hansen & Sutton, 2007) and the landscape patchiness in abiotic factors conducive to tree mortality (La Manna, Matteucci & Kitzberger, 2008). The variation in tree mortality of Austrocedrus chilensis has in turn effects on other forces affecting the vegetation, such as the browsing of an introduced deer species (Relva, Westerholm & Kitzberger, 2009).

Wherever the tree ring-width variability is poorly explained by variations in climatic parameters, other disturbance factors such as fire, pests, pathogens, and man are likely to have played an important role in determining tree health and growth in a certain landscape (Potito & MacDonald, 2008). Future scenarios of the impacts of climate change on forest health will need to take into account also the effects of climate change on these other forest disturbances. Although there is extensive evidence that climate change can profoundly affect tree mortality rates, few studies have established such an effect through the action of opportunistic fungal pathogens (but see Woods, Coates & Hamann, 2005; Marçais & Desprez-Loustau, 2007; Stone, Coop & Manter, 2008). Many more studies have dealt with the effects of climate change on tree mortality through changes in forest fire patterns, particularly in boreal forests and Mediterranean ecosystems.

By the end of the century, circumboreal forest fires are predicted to affect a doubled area, although this does not take into account that fire suppression efforts may be less effective in a warmer climate (Flannigan et al., 2009). Climate-induced forest fire changes may differ in different regions at the same altitude and latitude. For Russia, for example, climate change is predicted to increase the likelihood of forest fires in its European part and in Siberia, but not in the region around Lake Baikal (Malevsky-Malevich et al., 2008). However, previous, less sophisticated models suggested that a future general increase in fire activity in Russian and Canadian boreal forests is a virtual certainty (Stocks et al., 1998). Mismatches among the outcomes of models predicting the effect of climate change on forest fire behaviour can be explained not only by the different spatial resolution and parameterization of various models (Parisien & Moritz, 2009), but also by whether a gradual or an abrupt climate change scenario is considered. Rapid climate change is likely to be a determining factor for broad-scale fire activity (Marlon et al., 2009).

An increase in the number of days with potential fire danger has already happened for Siberia during the 20th Century (Groisman et al., 2007). A similar finding is reported for the Western USA, where the sudden increase in wildfire activity at the end of the 1980s has been linked to increased spring and summer temperatures (Westerling et al., 2006), rather than to a change in wildfire suppression policy. Since the 1980s, in association with warmer temperatures, stand-replacing fires have affected a larger area of forest in California and Western Nevada (Miller et al., 2009). For Canada, models at the beginning of the 1990s predicted that, in a world with doubled CO2 concentration, both the severity and the area burned could increase by one half (Flannigan & Van Wagner, 1991). In fact, seven of the nine years between 1997 and 2006 had extreme fires in Siberia, and years with an extreme forest fire situation have also become more frequent in Alaska and Canada (Soja et al., 2007). There is however a need for a more long-term perspective when assessing whether changes in forest fire activities can be related to changes in climate (Gillett et al., 2004; Girardin, 2007; Marlon et al., 2008).

By the end of the century, the increase in area burned under climate change in California is predicted to range between 9 and 15% above the historical norm (Lenihan et al., 2008a). Warmer temperatures are likely to increase fire risk through an increase in fuel flammability in currently relatively wet Californian forests (Westerling & Bryant, 2008). A return to the past drier conditions would lead to major changes in fire severity also in the boreal forests of North America and thus to shifts in tree species distributions (Fauria & Johnson, 2008). However, there is still much uncertainty about how climate change will affect wildfire intensity and frequency for single landscapes, given the many factors contributing to fire behaviour (Moritz & Stephens, 2008), the range in the forecasted levels of warming (Bachelet et al., 2001) and the trade-offs among different fire management strategies (Lenihan et al., 2008b). For world ecosystems, climate change during the 21st Century is predicted to increase the occurrence of wildfires in Amazonia, much of South America, many semiarid regions in Central Africa and Asia, across Australia and the Mediterranean, as well as in Northern regions such as Eastern Canada (Scholze et al., 2006).

Climate change is predicted to increase not only the susceptibility of forests to fire, but also to non-catastrophic disturbances such as fungal pathogens due to decreased growth and vitality of trees (e.g. for the mixed forests of California, Battles et al., 2008). A synergistic role will be played by storms, which are expected to increase in frequency and severity, thus leading to enhanced fuel quantities (Lindroth et al., 2009). The interactions between an increased forest fire activity due to climate change and other biotic (e.g. bark beetles: Jonsson et al., 2009) and abiotic disturbances will have repercussions on the carbon balance of forests. Increased forest fire, root rot, bark beetle and storm activity will make it likely that any negative feedback to climate change through increased carbon fixation due to higher CO2 concentration and longer growing seasons will be compensated by carbon release due to combustion and decay of organic matter (Field et al., 2007; Kurz et al., 2008). This is expected to be particularly the case in boreal forests, given that fire is already the major type of disturbance in these forests (e.g. Conard et al., 2002; Bond-Lamberty et al., 2007). However, forest fires are key contributors to carbon emissions in other biomes too. It is estimated that substituting wildfires with prescribed burning in the Mediterranean region would reduce carbon emissions by approximately half (Narayan et al., 2007), but it is unclear whether this result can be realistically achieved, given that changed climatic conditions are predicted to make Mediterranean landscapes even more fire-prone (Moriondo et al., 2006).

A reduction in carbon emissions can also be achieved with fire suppression, as documented for China between 1950 and 2000 (Lu et al., 2006). However, the North American experience of fire suppression (e.g. Hessburg & Agee, 2003; Donovan & Brown, 2007) casts doubt on whether this reduction can be maintained in the long term when landscapes will become more fire-prone due to the higher presence of flammable biomass and due to a warmer climate. Climate and forest fire regime changes have also an obvious impact on the conservation biology of (boreal) forests, where there is a dearth of initiatives to maintain (or establish) an effective network of protected areas in the face of the predicted effects of climate change (Martin, 1996; Li, Kräuchi & Gao, 2006; Scott & Lemieux, 2007). Climate change may also affect human health through increased forest fires, particularly in densely populated yet forested landscapes (Mestl et al., 2007; Kinney, 2008; Tham et al., 2009).

Climate change has the potential to affect landscapes throughout the tropics, particularly if these are already stressed by high levels of human disturbance (Verchot & Cooper, 2008), but also in the case of relatively undisturbed rain forests. The interactions of climate and land-use change with vegetation cover and plant health throughout the tropics will have important repercussions on the capacity of these ecosystems to sequester carbon (Kauffman, Hughes & Heider, 2009), as well as on their unmatched biodiversity. A four-year experimental drought (60% of the incoming precipitation was removed) increased large-tree mortality rates by a factor of 4.5 in an Amazonian forest (Nepstad et al., 2007). For Brazil as a whole, climate change is predicted to affect forest plantation productivity negatively mainly due to increased frequency and severity of droughts (Fearnside, 1999). Climate warming, if associated with an increased frequency of droughts, could well result in widespread forest fires for the Amazonian forest (Cochrane & Barber, 2009). This would provide a positive feedback on climate change by releasing further CO2 into the atmosphere (Golding & Betts, 2008). For the forests of the Kilimanjaro, climate warming is already believed to have contributed to the increase in forest fires, which, together with clearing for cultivation, have led to a reduction in cover (Hemp, 2009).

Since the mid-1970s, temperature has consistently increased in tropical regions in synchrony with the global trend (Malhi & Wright, 2004). At the same time, precipitation has decreased sharply in tropical Africa and marginally in tropical Asia, whereas no significant trend was detected for Amazonia, which is expected to be particularly susceptible to increased drought (Malhi & Wright, 2004). The occurrence of a multi-year drought has the potential to counteract any increases in tree cover of xeric savannahs due to previous wet periods, as shown in Australia by Fensham, Fairfax & Ward (2009), with consequences for carbon storage in these ecosystems. In addition to changes in precipitation patterns, the future health of tropical forests will be heavily influenced by the synergisms between logging and forest fires (Siegert et al., 2001; Cochrane & Laurance, 2008). There is a need for studies of long-term time series to unravel which socio-economic factors can explain local and regional variation in deforestation rates (Ewers, Laurance & Souza, 2008), but we also need to make sure that the forecasted near-term dieback of vast parts of the Amazonian forest is avoided thanks to the establishment of effective protected areas, fire control, and carbon market incentives (Nepstad et al., 2008), which can also contribute to a more sustainable development of the region (Gardner et al., 2009; Rodrigues et al., 2009).

One factor which is likely to affect (sub)tropical forested landscapes, their carbon sequestration capacity and their management in the near future, and which can jeopardize efforts at achieving sustainability, is the predicted increased frequency and severity of tropical hurricanes (Stanturf, Goodrick & Outcalt, 2007; Hopkinson et al., 2008; Laurance & Curran, 2008; Lugo, 2008; Zeng et al., 2009). This will be particularly the case for forests in coastal regions, and will thus pertain especially to mangrove forests. Mangrove forests, which occur throughout the tropics, have been historically resilient to changes in sea-level rise, but are currently threatened by deforestation and other human disturbances. This makes them less at risk from future climate change than from other current anthropogenic activities (Alongi, 2008). However, sea-level rise may yet prove to be the greatest long-term threat to mangrove forests, especially where there is little room for landward migration (Gilman et al., 2008). The decline of coastal forests along the Gulf Coast in Florida has been shown to be accelerated by the interaction of sea-level rise and increased drought (Desantis et al., 2007).

Some general patterns predicted for temperate landscapes are likely to be appropriate also for the tropics. For example, shifts in tree species distributions due to climate change are expected also for (sub)tropical ecosystems (e.g. Enquist, 2002). One example is provided by a population census throughout the distributional range of Aloe dichotoma, a Namib desert tree, which was shown to be retreating at the equatorward range edge but not expanding at the poleward range edge (Foden et al., 2007). There is a need to assess how general such asymmetric processes are for other tree species in the tropics (Delire, Ngomanda & Jolly, 2008; Platts et al., 2008). As for tree land cover, the issue of tropical tree species distribution shifts in response to climate change will be confounded by the varying intensity of land use in various regions (Higgins, 2007). As for temperate and boreal tree species, but all the more so given the often restricted ranges of tropical tree species, genetic diversity is expected to play a major role also in the adaptability of tropical trees to changed environmental conditions (Bawa & Dayanandan, 1998).

Recent genetic studies are providing evidence that tree species may have survived glaciation events in northerly refugia (e.g. Picea glauca in Alaska; Anderson et al., 2006). Such findings seem to be general at least for tree species with small seed size (Bhagwat & Willis, 2008; Birks & Willis, 2008). They imply that the postglacial migration rate of tree species calculated assuming survival exclusively in southern refugia was probably overestimated. If such previously unknown glacial refugia were common for many species (Stewart & Lister, 2001), the capacity of tree species to react to future climate warming by migrating towards the poles might be lower than previously thought. Although investigations of the effects of past climate changes on tree genetic diversity can in some cases provide lessons for the future, past data can hardly be used to validate models of future changes in tree species distribution given the unprecedented speed of the predicted changes in climate and the pervasiveness of other human influences on the current landscape (Hampe & Petit, 2005; Petit, Hampe & Cheddadi, 2005; Petit, Hu & Dick, 2008).

In spite of the plasticity inherent in tree species phenotypes (e.g. Mboumba & Ward, 2008), there is some evidence that future climate change may bring about a pervasive maladaptation of tree species, although care needs to be taken not to mix conclusions and predictions based on neutral and adaptive genetic variation (Kozlowski & Pallardy, 2002). An example of this approach is a study of seedlings of coastal Pseudotsuga menziesii (var. menziesii), which were shown to be at high risk for most traits particularly in the case of the stronger climate change scenarios (St Clair & Howe, 2007). Another example is a study of seedlings of Fraxinus americana, whose Eastern populations appear not to be adapted to the drier conditions forecast in the Eastern USA (Marchin, Sage & Ward, 2008). A similar sobering finding is the evidence for southern bottlenecks in the range of Picea engelmanni, which suggests that this species is particularly threatened by future global warming (Ledig, Hodgskiss & Johnson, 2006). More studies are needed of the association of the genetic diversity of tree species and their associated organisms (e.g. rhizobes in N-fixing tree species, endophytes, mycorrhiza) with resistance to drought, salinity, increased temperature and other stresses.

Tree species genetic diversity in mountainous regions appears to be higher at intermediate altitudes, although exceptions to this pattern are present (Ohsawa & Ide, 2008). This finding is of relevance for future climate change, as low-lying regions are predicted to become suboptimal for many species, and as tree populations confined to mountain summits by their adaptation to cold climates are likely to dwindle or disappear (Ohsawa & Ide, 2008). Since tree populations at intermediate altitudes are thus likely to play an important role in the reaction of these species to climate warming, their higher genetic diversity is to be welcomed, as it provides more resilience in the face of environmental change. For Fagus sylvatica, mountains appear not to have posed a barrier to recolonization following glaciation (although the Alps coincide with a genetic discontinuity between Italian and other beech populations), but to have facilitated its spread towards northern latitudes, which was instead hindered by plains such as the Danube valley (Magri et al., 2006; Magri, 2008). Mountains and plateaus may play a similar role in future migrations.

The unprecedented rates of climate change, coupled with land-use patterns which impede rapid tree migration, are expected to pose a substantial threat to many tree species (Davis & Shaw, 2001; Parmesan, 2006; Mckenney et al., 2007). This threat will be worsened by the often lower genetic diversity showed by tree populations at the margins of their distributional range, given that migration to track climate change will tend to start from these boundary populations (Savolainen, Pyhäjärvi & Knürr, 2007). If climate change acts in the presence of low genetic variation, the ability of a species to respond to the changing climate is potentially lower (Hellmann & Pineda-Krch, 2007). However, low genetic diversity in range-edge populations could result from selection for particular alleles that make these populations adapted to conditions at the range margin. Should this be the case, low genetic variation might not manifestly impair the response to climate change of these peripheral populations.

Similar studies of the adaptive genetic diversity of grassland species are needed to assess how resilient such ecosystems will be to changes in climatic conditions (e.g. Knapp & Rice, 1996; Kolliker et al., 1999; Casler et al., 2007; Rudmann-Maurer et al., 2007; Bylebyl, Poschlod & Reisch, 2008). More than with forest trees, studies of the geographical genetics of grassland species can be made difficult by past movement of propagating material, particularly for grassland species commercially grown in managed grassland. Not all currently existing grasslands are agricultural, but at least in Europe there has been a strong reduction in the area and connectivity of semi-natural grasslands in the last century (Honnay et al., 2007). Agriculture is now the major use of land globally with some 1.2–1.5 billion hectares under crops, 3.5 billion hectares grazed and another 4 billion hectares of forest used by humans to varying extents (Howden et al., 2007). Grasslands cover 40% of the earth's land surface and provide multi-use ecosystem services, including animal-based industries, maintenance of soil cover and biodiversity, sequestration and storage of atmospheric CO2, and have tourism and leisure value (Chakraborty, 2001). Plant disease decreases herbage productivity and palatability as well as seed yield although quantitative data compared with arable and horticultural landscapes are rare. There are also effects on species composition and thus on the other services mentioned above.

Pathogens will interact with climate change to affect the biodiversity and sustainability of grassland ecosystems (Roy, Güsewell & Harte, 2004). The future risk of exotic and invasive pathogens will thus provide a major challenge to grassland health. There is a strong rationale for an increased focus on adaptation of agriculture to global climate change. Such adaptation may provide opportunities for agricultural investment (Howden et al., 2007). Grassland agriculture has also been proposed as potentially providing some mitigation options for climate-change effects on agriculture (Hopkins, 2003). These options include the growing of biofuel crops to displace carbon fuels, managing carbon sequestration by grassland soil and vegetation, and measures to reduce other greenhouse gas emissions. However, the climate-change mitigation contribution of biofuels is even more disputed when these are compared to permanent unimproved grassland (Rowe, Street & Taylor, 2009). Moreover, the landscape-scale adoption of the grass Miscanthus spp. and similar energy grasses instead of short rotation coppices and semi-natural woodland may have unintended consequences for biodiversity and hydrology (Rowe et al., 2009).

The changing global climate presents a major challenge for grassland ecology and more applied aspects such as grass breeding (Humphreys et al., 2005; Tuba & Kaligariè, 2008). Global warming will affect temperature-sensitive traits such as vernalisation. Grass biomass use may contribute to reductions in greenhouse gases emissions, but also in an increase due to methane production by cattle. Because of the unpredictable outcomes of climate change, grass breeders will need to have a range of germplasm available for new applications. The paucity of long-term data sets relating to grassland landscapes, both natural and managed for grazing, makes the prediction of the effects of climate change on productivity and biodiversity difficult. Data on grassland plant community structures at the Park Grass experiments, Rothamsted Research, UK, have been collected over a 90 year period (Leigh & Johnston, 1994). These data were used to analyse composition in relation to annual biomass and rainfall variation (Silvertown et al., 1994). Although the statistical measures gave conflicting results, composition was related to biomass variation, itself related to rainfall variation. Biomass was significantly increased by rainfall which selectively favoured grasses. Asymmetric competition magnified the effect of rainfall on community composition. Long-term experiments in agricultural and grassland soils are also invaluable for predicting the effects of global warming on C-decomposition in soil, and hence CO2 release. Models developed at Rothamsted Research indicate that treating the top layers of soil as a homogeneous unit can greatly overestimate carbon release due to an increase in temperature (Jenkinson & Coleman, 2008). There is a potential link with plant health, as carbon decomposition in the soil provides nutrients to plants, so that if climate warming makes this decomposition faster, some plants may be favoured in the short term, although in the long term the fertility from the soil may be depleted.

The importance of biotic stress factors, such as weed, insect pest and disease interactions with climate change is understood qualitatively but quantitative knowledge is lacking (Tubiello et al., 2007), certainly for pastures. Effects of climate change on grassland systems are manifested at different levels of integration from the cell to the sward to the landscape. There are problems in integrating responses between different levels (Booth & Grime, 2003). Although responses at the lower plant level have been described with some confidence, higher level interactions with soil biota, pests, weeds and diseases are still poorly understood and may be critical in determining the response of the whole system (Pollock et al., 1995).

Experimental reductions in plant richness increased the vulnerability of grassland ecosystems to invasive plant species, facilitated the spread of fungal plant pathogens, and altered the structure of insect communities (Knops et al., 1999). This study shows that plant health can be a function of ecosystem structure, which is in turn connected with other services provided by plant assemblages. Similarly, the disease severity of rust fungi on Lolium perenne decreased with increasing species richness of experimental grass communities (Roscher et al., 2007). These findings support the ecological theory that loss of biodiversity and species at a basal level impacts on entire ecosystem functioning. In a study involving 24 grassland plant species and 11 diseases, grassland communities with reductions in the number of less-disease-prone species had higher pathogen loads; a similar result was fond but to a lesser extent when communities lost the species dominant at high diversity (Mitchell, Tilman & Groth, 2002). These results support the hypothesis that decreased species diversity increases foliar pathogen load if disease-prone host abundance and therefore disease transmission is increased. Furthermore, elevated CO2 and nitrogen addition increased the pathogen load (Mitchell et al., 2003), suggesting that an increase in pathogen load can be an important mechanism by which global change affects grassland ecosystems. Interestingly, there are now short-term experiments on the effects of environmental change in natural populations such as prairie species, at the gene expression level (Garrett et al., 2006). Gene expression associated with a hypersensitive response (such as occurs with pathogen challenge) in Andropogon gerardii under different precipitation patterns suggests a significant defensive cost associated with climate change (Travers et al., 2007).

There are very few long-term data sets of plant pathogen incidence in grassland species. One good example is a study of Barley yellow dwarf viruses preserved in herbarium specimens of California grasses (Malmstrom et al., 2007) from 1917. The herbarium evidence and phylogenetic analysis of sequences suggest that the viruses were present in native grasses at the time they were being displaced by Eurasian grasses, indeed facilitating invasion at the landscape level. Strains of Barley yellow dwarf virus (BYDV) infect many of the common grass species present in a tallgrass prairie in Kansas (Garrett et al., 2004). The effects of BYDV on prairie plant communities remain to be determined. The incidence of fire in these prairie communities, and indeed in perennial grain species such as wheatgrass (Thinopyrum spp.), may decrease disease pressure by decreasing inoculum sources (Cox, Garrett & Bockus, 2005). This example shows that many of the ways through which global change will affect plant health (in this case plant disease, invasive species and wildfire regimes) can co-occur in the same situation. Hence the need for considering all these factors in any serious assessment of future plant health under global change.

There is now a consensus that climate change will not only affect landscapes by disrupting the historical regime in climate and disturbances, but also by acting synergistically with other artificial processes such as an increase in trade (Jones & Baker, 2007; Sutherst, Maywald & Bourne, 2007b; Ward & Masters, 2007; Sutherland et al., 2008; Waage & Mumford, 2008). Together with an increasing globalisation of economic activities, climate change is now recognized to be one of the main drivers accelerating biological invasions (Perrings et al., 2005; Hobbs & Mooney, 2005; Fig. 2). Given that plant pathosystems are composed of the interaction between hosts and pathogens (in a disease-prone environment), also hosts are key for an understanding of pathosystems. This justifies dealing with plant invasive species in a review of the effects of global change on plant health. Moreover, if a plant species invades a new region, and if plant pathogens accompany this invasion (or if pathogens already present are able to infect the new host), a new pathosystem will be formed, which would not have been present without the plant invasion.

Figure 2. Disease triangles (composed by the interactions of host, pathogen, and environment) under global change (climate change together with other worldwide anthropogenic processes such as increasing land use change, air pollution and intercontinental trade); arrows are symbols for the long-distance movement of pathogens enabled by increased globalization, dashed triangles represent pathosystems which are not noticed by disease managers but nonetheless contribute to the epidemic development (modified from Jeger & Pautasso, 2008).

Climate change is expected to have different impacts on different stages of the invasion process (Theoharides & Dukes, 2007; Hellmann et al., 2008). At the first stage of this process, the uptake of species in their native area and subsequent transport, the chances of species survival during transport may, for example, increase due to shorter shipping times caused by the loss of Arctic sea ice (Hellmann et al., 2008). However, the effect of increasing trade is likely to be far larger (Perrings et al., 2005; Dehnen-Schmutz et al., 2007; Meyerson & Mooney, 2007). Using historical data on merchandise import volumes and non-native species arrival rates in the US, Levine & D’Antonio (2003) modelled the expected numbers of future invaders. For plant pathogens, they forecast 5 to 61 new arrivals by 2020 depending on the modelling approach used. Interceptions of non-indigenous plant pests and pathogens at 160 US border inspection points increased from around 20,000 in 1984 to more than 50,000 in 2000 (McCullough et al., 2006). These interceptions represented at least 2,340 species coming from 259 different geographical origins.

At the arrival and establishment stage of the invasion, climatic conditions (together with competition, propagule density, herbivores and pathogens) in the recipient region determine whether or not an introduced species will survive and establish in the new environment. Climatic conditions are therefore an important criterion used in the analysis of species invasion success, risk assessment procedures designed to identify future invaders (e.g. Pheloung, Williams & Halloy, 1999; Rayment, 2006; Keller, Lodge & Finnoff, 2007), and the modelling of invasive species distributions (e.g. Sutherst & Maywald, 1985; Thuiller et al., 2005). Predicted climate change has added a new dimension to this stage. Climate change will not just allow the survival of species that would not have been adapted to the previous climate. It will also enable a new range of interactions resulting from increasing habitat disturbances, shifts in the range of native species, and the increased load of novel pathogens (Crowl et al., 2008). Furthermore, species that may have passed a risk assessment procedure because the risk of their establishment in unintended habitats had been considered very low due to the non-matching of climate variables may now be present in countries where they could become problematic invaders in the future.

Similarly, species already present in a new area but not able to establish or spread may do so under novel climatic conditions. This may be particularly the case for ornamental plants where a vast pool of alien species is already present in gardens. Gardens have already been the source of the majority of alien plants established in many floras. An example is provided by naturalization of the Asian palm Trachycarpus fortunei in southern Switzerland starting from plantations in gardens and parks from the 1950s onwards. The establishment of this species has been linked to an increased mean of monthly January temperatures and a continuous northwards shift of suitable habitat conditions (Walther et al., 2007). The expansion of tree mallow (Lavatera arborea) on the Scottish island of Craigleith from less than 5% land cover before 1960 to 85% today has been correlated to a rise in winter temperature (van der Wal et al., 2008). This case study also underlines the importance of other anthropogenic factors in such expansions. In this case, important factors were the decline in the island's rabbit population (due to the introduction of a disease) and the increase of soil nutrients (following an expansion of the puffin population as a result of changing fishing practices). Adaptation strategies to climate change may also involve the introduction and large-scale plantation of species previously not present in an area as in the case of biofuels where suitable species often exhibit traits that have been shown to favour species invasion (Raghu et al., 2006; Barney & DiTomaso, 2008). The example of biofuels shows that climate-change-mitigation options can in turn raise new problems, so that an integrated analysis of potential effects is recommended.

Climate is one of the main determinants of species distributions. Climate-change effects on native species distributions have thus often been modelled. Similar modelling has recently being used to forecast the potential distribution of invasive species. An example is the use of the CLIMEX modelling software (originally developed to predict species distributions for the purpose of quarantine, biological control, pest management and epidemiology; Sutherst & Maywald, 1985) to predict the potential change of distribution of the invasive species Acacia nilotica ssp. indica in Australia (Kriticos et al., 2003). Together with predicted changes in temperature, the model also includes the expected higher water-use efficiency under increased atmospheric CO2 concentrations enabling A. nilotica to extend its range into previously unfavourable habitats. At a different scale, and using ecological niche models, Peterson et al. (2008) analyse the invasive potential of around 1800 European plant species. The models predict an overall decline in areas outside of Europe suitable for invasion, because many of these European plant species are limited in their distribution by warm climates.

With a different approach, Thomas & Ohlemüller (2010) derive a global map of invasibility by calculating an invasibility index. This is the ratio of similar climates within and outside a distance of 1000 km from the analysed 0.5° grid cell, and is based on the assumption that large areas with similar climates outside this range may harbour a large pool of species able to establish. Different climate-change scenarios are then used to map areas with increasing and decreasing invasion risks. Such general approaches offer the potential for a rapid screening procedure for large species datasets or the identification of high risk areas. However, it is unclear how precise these predictions are likely to be over the long term (Sax et al., 2007; Jeschke & Strayer, 2008).

The question arises therefore if current prevention and management measures against invasive species will be sufficient under climate-change scenarios. There are already many limitations in our landscape management of invasive species without taking climate change into account (Simberloff, Parker & Windle, 2005; Hulme et al., 2008). Conversely, even in the presence of climate change, land use and human-mediated dispersal will still be important forces driving the invasion of exotic species (Hulme, 2009). Analysing the movement and introduction of forest pathogens through increasing horticultural trade, Brasier (2008) concludes that the current biosecurity regulations are today insufficient to prevent future landscape outbreaks of serious diseases. A paramount example is provided by Phytophthora ramorum, the oomycete responsible for Sudden Oak Death in California and one county in Oregon, which is thought to originate from South-East Asia (e.g. Grünwald, Goss & Press, 2008; Fig. 3). Brasier (2008) argues that invasive species are causing increasing damage because the current biosecurity framework was developed in the 1950s, a time of much less intercontinental trade. There is certainly a need to include in biosecurity regulations not only the changed global environment, but also the increased availability of data on phytosanitary inspections of imported plants (Surkov et al., 2008) and recent advancements in network theory applied to the spread of diseases and invasive species (Jeger et al., 2007; Harwood et al., 2009; Moslonka-Lefebvre, Pautasso & Jeger, 2009; Pellis, Ferguson & Fraser, 2009).

Here we provide a summary of the broad issues which will confront plant health managers under global change. Selected practical recommendations for landscape managers are provided in the conclusions. These headlines might not be relevant for all specific conditions and will need to be integrated with local knowledge, successful approaches hidden in the grey literature and the historical legacies of past management. There are three main strategies to safeguard plant health under global change: (i) adaptive conservation, (ii) assisted migration and (iii) enhancing diversity. A diversity of adopted strategies (including doing nothing) can itself be a good thing, so as to minimize the likelihood of committing mistakes over large regions. In the long term, it will be fundamental to include the issues treated in this and other reviews in the curriculum of schools and universities, and not just for students of plant pathology, so as to foster awareness, dissemination, interdisciplinarity, as well as innovation. For example, it is time to scale up from agri-environment schemes focusing on individual fields and farms, to an approach at the landscape level, with targeted incentives for regions of particular importance e.g. for preserving levels of genetic and species diversity in warmer and/or drier conditions.

A key point for landscape management of invasive species under global change will be the changing perception of what is regarded as alien and native in a changing climate. Thomas & Ohlemüller (2010) point out that some species will only be able to survive in regions where they would be regarded as alien with the distribution of others expanding increasingly into new regions. The number and extent of novel ecosystems (Hobbs et al., 2006) consisting of species compositions previously not seen and resulting from human activities will therefore increase and conservation policies will need to take that into account if global biodiversity is to be preserved.

There are three main strategies for living organisms to withstand climate change: (i) by use of their phenotypic plasticity, (ii) by evolving new traits, and (iii) by migrating to areas more suitable to their physiological requirements, or by a combination of these three strategies (St Clair & Howe, 2007; Jump et al., 2008; Nitschke & Innes, 2008). Although disentangling micro-evolutionary adaptations to climate change from plastic phenotypic responses is revealing itself particularly challenging (Gienapp et al., 2008), only models of species vulnerability to climate change which take into account these three strategies are likely to deliver realistic decision-support tools for landscape managers (Davi et al., 2006; Reusch & Wood, 2007; Xu, Gertner & Scheller, 2007).

Landscapes which have been historically already subject to stress and climatic variation have particular conservation value because they may contain genetic adaptation to climatic extremes (Gitlin et al., 2006). But in ecosystems such as those along the Mediterranean Sea, where human activities have caused ecosystem stress for millennia, landscape fragmentation will make it unlikely that genetic variation of tree species will be preserved in the few patches of semi-natural forest left (de Dios, Fischer & Colinas, 2007). Although it has been shown that plant diversity tracked past variations in climate without any assistance from man (Willis et al., 2007b), the unprecedented speed of the forecasted changes has kindled a debate in conservation biology about assisted migration (e.g. Hunter, 2007; McLachlan, Hellmann & Schwartz, 2007; Hannah, 2008; Hoegh-Guldberg et al., 2008; Ricciardi & Simberloff, 2009; Richardson et al., 2009). Assisted migration consists in facilitating long-distance dispersal of organisms which would be unable to migrate fast enough to track climate change. Assisted migration will thus be of relevance for species with small and fragmented populations, and for those debilitated by diseases and abiotic disturbances (Aitken et al., 2008). Studies of the landscape genetics of organisms threatened by climate change should inform any future efforts at assisting their migration, because the source of propagule material can determine the outcome of these activities depending on the amount of adaptive genetic variation in this material. Similar investigations, however, need to be pursued for associated organisms such as pathogens, endophytes and mychorrhiza, as these are also likely to be disseminated through the assisted migration of their hosts (Brenn et al., 2008; Vellinga, Wolfe & Pringle, 2009).

Given that the distributional ranges of species are shifting and will continue to shift in response to climate change, current protected area networks have become or are likely to become outdated. It will thus be necessary to keep track of biodiversity hotspots and assist also the migration of conservation activities (Hannah et al., 2007; Normand, Svenning & Skov, 2007; Pressey et al., 2007; Hole et al., 2009). How to preserve old-growth forests in this dynamic situation remains an outstanding challenge (Bauhus, Puettmann & Messier, 2009). However, climate change may provide an opportunity to increase the relevance of protected areas for conservation, as in many countries these have historically been located in landscapes of low productivity, low human population density, and low species richness (e.g. Luck, 2007; Virkkala & Rajasarkka, 2007; Loucks et al., 2008). Many recent studies have indeed shown a current large-scale coincidence of species-rich and densely inhabited regions (Balmford et al., 2001; Araújo, 2003; Pautasso, 2007; Marini et al., 2008). This has obvious implications for current conservation efforts, as human beings generally have detrimental impacts on biodiversity at a local scale and these effects are compounded if, over regional to continental scales, people have preferentially settled and become more numerous where there are more species. It would be interesting to know in advance how this human population-biodiversity correlation over large scales will develop in the presence of global warming: provided that the current fossil fuel bonanza and related human mobility is sustained (Diamond, 1994; Schulz, 2004; Haber, 2007), it is likely that people will be much quicker than other species in migrating out of drier or flooded conditions. Similarly, it would be timely to extend studies of the regional species-people correlation from analysis of well-known taxa such as birds and mammals to a less investigated part of biodiversity such as plant pathogens.

For landscapes where tree mortality patterns have been artificially lowered by fire suppression, it would be theoretically important to reintroduce forest fires in the range of sizes and frequencies known to have occurred historically. Practically, however, decade-long fire suppression has made it very difficult for prescribed burning to mimic the natural fire regime, and climate warming and the growing forest-urban interface will make it even more unlikely that such a regime could be achieved (Syphard et al., 2007; Chapin et al., 2008; Fulé, 2008). Indeed, models predict that the number of fires escaping control in California could increase due to climate change, although an increase in firefighting resources might still initially be able to compensate for this increased risk (Fried et al., 2008). Wildfire under climate change exemplifies the extent to which landscape managers may be underestimating future challenges to established practices. For California, land and fire managers appear to have limited awareness of the threat posed by climate change in terms of increased likelihood of uncontrolled fires, which translates into a limited potential capacity to withstand this increased threat, and equally limited practical actions undertaken so far to solve this issue (Moser & Luers, 2008). A similar argument can be made for plant diseases and invasive plants in many other regions predicted to experience a shift in climatic mean and extreme values.

Future plant health will be affected by climate change through a number of landscape-scale processes, from changes in mortality rates of long-lived species such as trees to the effects of invasive species. Wildfires, bark beetles, droughts, new diseases and shifts in tree species distributions will all contribute to push forest ecosystems out of their familiar dynamic. In these concluding remarks, we draw together some key messages from the reviewed body of literature. We provide a reminder of the many factors affecting plant health under a changed climate which will need to be considered when planning future research on plant health. Much difficulty in predicting future developments will be due to the potential interactions among these factors. Stochasticity, internal model variability, and the likely rapid evolution of both plant hosts and pathogens in the presence of climate change are additional hurdles for the community of modelers. The contribution of various plant diseases on different host species and the long-term nature of climate change are only some examples of the research challenges before us.

A further difficulty stems from internal climate model variability. This has been shown to be able to lead to substantial differences in the projected changes of future distributions of species in response to climate change (Beaumont et al., 2007). It is likely that also for plant pathosystems and related abiotic disturbances, the differences among different realizations of a single model predicting effects of climate change can be greater than differences resulting from the variability among different models. Much further complexity can be added to these scenarios by considering that not only plant host species (e.g. Jump & Peñuelas, 2005; Morin & Thuiller, 2009), but also plant pathogens will show evolutionary adaptability to climate change (e.g. Scholthof, 2007; Lebarbenchon et al., 2008; Mace & Purvis, 2008; Whitney & Gabler, 2008). However, if changes in abiotic conditions, e.g. air quality (Wang et al., 2007), are rapid, there will probably not be enough time for evolution to act, neither for hosts nor for pathogens.

Most models of future impacts of climate change on plant diseases have focused on single pathosystems (e.g. Brasier & Scott, 1994; Bergot et al., 2004; Evans et al., 2008; Dukes et al., 2009). But the future health of natural ecosystems will be the outcome not only of single interactions between pathogens and their hosts, but of all the relevant pathosystems summed up in a certain habitat type (Walther, 2004; Mulder, Roy & Güsewell, 2008; Vacher et al., 2008). There are exceptions in experimental studies. For example, in the grassland-warming experiment performed in Colorado, USA, the relative abundance of several plant pathogens was assessed (Roy et al., 2004). Simulations of the long-term development of plant pathosystems need to integrate the inherent diversity of natural plant communities and of their pathogens and the effects that this diversity may have on the resilience of the vegetation.

Climate change is a worldwide, long-term process. The implications of an increase in global temperature of up to 4°C by 2100 and of associated changes in precipitation for plant health, land use and the rural economy need thus to be studied with long-term and broad-scale data sets and in a continuous, adaptive and international research framework (Kerr, Kharouba & Currie, 2007; Thuiller, 2007; Willis et al., 2007a; Garrett, 2008; Marshall et al., 2008; Woodward & Kelly, 2008). Some examples of long-term research on the effects of global change on plant health are appearing, from plant health studies using data from the long-term Rothamsted fertilizer experiment (e.g. Poulton, 1996; Bearchell et al., 2005; Ogilvie, Hirsch & Johnston, 2008; Shaw et al., 2008) to analyses of the incidence of Phytophthora infestans in relation to climate parameters in the Netherlands and Finland (Zwankhuizen & Zadoks, 2002; Hannukkala et al., 2007), but much of this research is still performed over a restricted spatial scale. A long-term perspective is needed also for recommendations for landscape planning. For example, if an association between drought and tree mortality can only be detected from multi-year periods (as, for example, in old-growth mixed conifer forests in Yosemite National Park; Guarin & Taylor, 2005), then it is important for managers to consider long-term trends in climatic variability. Similarly, a regional and not just local perspective in adapting land use to climate change is fundamental, as alterations of ecosystem type, associated ecosystem properties, and land surface conditions will be particularly marked for regional-scale tree mortality and migration (e.g. Orwig, 2002; Breshears et al., 2005; Gray et al., 2006; Fischer et al., 2009).

There are a number of research gaps and interdisciplinary opportunities for plant epidemiologists in the coming decades: (i) the integration of short-term, local experiments with landscape-scale, long-term modeling of plant health, (ii) the merging of scenarios of the effects of predicted climate and land-use changes on both host and pathogen species, and (iii) the use of insights and techniques available from studies of the consequences of global change for animal and human diseases. Future challenges in this rapidly growing field will include spatial and temporal scale-dependences in the shifts of the environment and of plant pathosystems, co-evolutionary interactions between hosts and pathogens, the use of plant pathogens as indicators of climate change (Garrett et al., 2009) and signal accumulations and lag effects not only in the responses of pathosystems but also in the ability of landscape managers to adapt to changing conditions.

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